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195
Assessing the Impact of Different Redox Conditions and Residence Times on the
Fate of Organic Micropollutants during Riverbank Filtration
Carsten K. Schmidt, Ph.D.;
Frank Thomas Lange and Heinz-Jürgen Brauch,
DVGW-Technologiezentrum Wasser (TZW), Karlsruher Str. 84, 76193 Karlsruhe, Germany
Abstract
Riverbank filtration and artificial groundwater recharge are well established techniques in Germany that are
often used as an important component of the established multiple barrier system. Passage of water
underground provides several benefits for drinking water treatment. Experience demonstrates that during
infiltration and underground transport, processes such as filtration, sorption, and biodegradation produce
significant improvements in raw water quality. However, due to industrial and municipal discharges and the
influence of agriculture, rivers and lakes are polluted by a number of organic chemicals. In order to assess
the impact of those organic micropollutants on the quality of drinking water it is necessary to clarify their
fate during infiltration and underground passage.
The fate of organic micropollutants in a river water-groundwater infiltration system is mainly determined
by adsorption mechanisms and biological transformations. The purification process of the underground
passage starts in the infiltration zone that can be characterized as a gelatinous, biological highly active
biolayer which consists of algae, bacteria, fungi and protozoa as well as of organic and inorganic particles.
During infiltration, the river water with its dissolved components meets multifaceted biogenous and
abiogenous surface structures that aid in intensifying most of the self-cleaning mechanisms that are in
principle also present in the free surface water. During biodegradation of substances in water, dissolved and
chemically fixed oxygen is consumed, thereby causing shifts in the underlying redox system that affect
form and extent of microbial degradation processes. In these redox reactions, available organic substances
represent electron donators, while reducible substances in water (O2, NO3
-, NO2
-, SO4
2-) and solid phases
(Fe(III)-, Mn(IV)-oxides and -hydroxides) act as electron acceptors. In an anaerobic aquifer, concentrations
of oxygen, nitrate and sulfate are subsequently depleted whereas levels of ammonia and sulfides increase.
At the end of this sequence, that is possibly formed already in the first centimeters of the infiltration zone,
stands the reduction of carbon dioxide with the corresponding formation of methane. Depending on local
conditions, individual redox zones may vary considerably with regard to their spatial expansion. Since
many microorganisms can perform their microbial activity only at certain redox potentials and since the
extent of biodegradation is dependent on sufficient residence times and sometimes proceeds retarded, the
degradation of organic micropollutants is linked to both the presence of favorable redox conditions and
sufficient residence times in these zones. As a consequence, elimination rates of intrinsically degradable
substances may vary considerably depending on local geological and hydrochemical conditions as well as
on organic loads of surface waters and infiltration zones.
For investigation of the impact of different boundary conditions on the purification capacity of bank
filtration, extensive measurement campaigns were carried out at four well-characterized aquifers along the
rivers Rhine, Ruhr and Elbe. The different characteristics of these infiltration pathways allowed for detailed
investigations concerning the elimination capacity of riverbank filtration at various redox conditions
(aerobic, aerobic-denitrifying, denitrifying, strictly anaerobic) and residence times of the infiltrated water in
the underground (5-300 days). Analyzed target compounds comprised several contaminants relevant for the
aquatic environment, such as complexing agents, aromatic sulfonates, pharmaceuticals (including iodinated
X-ray contrast media), and MTBE.
On the basis of a comprehensive evaluation of these investigations, it is obvious that the removal efficiency
of bank filtration for different polar organic micropollutants is extremely dependent on the underlying redox
processes in the aquifer, since some organic micropollutants (e.g. iodinated X-ray contrast media,
carbamazepine, sulfamethoxazol) turned out to be better removable during denitrifying and strictly
anaerobic bank filtration, while others (e.g. EDTA, DTPA, diclofenac) were better degradable in aerobic
aquifers. Furthermore, it is quite difficult to predict the fate of single substances during bank filtration, since
similar compounds often demonstrate huge differences in their elimination rates. The elimination takes
place predominantly in the first few meters of the infiltration pathway, even though longer residence times
can significantly improve removal rates of individual micropollutants (e.g. naphthalene-1,3,6-trisulfonate).
196
Introduction
In Germany, groundwater is used for drinking water production wherever possible. When compared with
surface water, groundwater is well protected against most types of pollution, is of relatively regular quality
and temperature, and its abstraction can be easily adjusted to short-term fluctuations in consumption.
However, exploitation of natural groundwater sources is restricted with regard to quantity. In Germany, this
limitation is not given to such an extent for surface waters. However, surface water, particularly river water,
is exposed to dangers of permanent and sudden pollution by wastewaters or to disturbances due to storage,
transport or application of water-endangering substances, thereby always reflecting its function as receiving
water. In order to preserve the protective character of groundwater at least partly when utilizing surface
water for drinking water preparation, surface water is subjected to an underground passage via bank
filtration or artificial groundwater recharge. Experience demonstrates that during infiltration underground
transport processes such as filtration, sorption, and biodegradation produce significant improvements in raw
water quality. Underground passage as water treatment procedure combines particle removal, pathogen
removal, organic and inorganic chemical removal, peak smoothing in spills, temperature equalization,
reduction in DBP formation, and production of a more biologically stable water [1-3].
The effectiveness of bank filtration and artificial groundwater recharge has long been recognized in
Germany. As a consequence of various bacterial diseases caused by drinking water from waterworks with
direct intake from rivers in the late 19th century (e.g. outbreak of epidemic cholera in Hamburg in 1892/93),
direct extraction of surface water for public water supply fell into discredit and was replaced or
supplemented by artificial or natural subsoil passage of river water due to its efficiency in removing
microorganisms from the infiltrating surface water. Nowadays, approximately 16 % of the drinking water in
Germany is produced from bank filtrate or infiltrate. Because of pollution, direct treatment of river water
has dropped to 1 %. Water suppliers in Berlin produce approximately 75 % of the drinking water by bank
filtration and artificial groundwater recharge. In Germany, more than 300 water works use bank filtration
and roughly 50 plants are based on artificial groundwater recharge [1].
Surface water is often affected by industrial, agricultural, and domestic pollution. Various organic
micropollutants have been detected in surface waters. Polar organic molecules, such as complexing agents,
pesticides, industrial products like aromatic sulfonates, pharmaceutical compounds, and personal care
products, are of recent concern. Bank filtration can significantly lower the concentrations of many surface-
water pollutants. However, precise predicting and quantifying such reductions is often difficult, since the
efficiency of the underground passage depends on several factors. These include the river water quality,
geological conditions, porosity of the soil, residence time of the water in the soil, temperature, pH-
conditions, and redox status. Thus, the behavior of chemicals and microorganisms during infiltration and
underground passage of water depends on many different interacting factors.
The infiltration of surface water into the aquifer is a filtration process in which complex physical, chemical
and biological factors have a combined effect on the purification of the infiltrated water. The purification
process starts in the infiltration zone that can be characterized as a gelatinous, biologically highly active
biolayer which consists of algae, fungi, protozoa as well as of organic and inorganic particles. Purification
processes taking place during infiltration are similar to the self-cleaning properties found in surface waters
but proceed in the infiltration zone much more intensively. The fate of organic micropollutants is mainly
determined by adsorption mechanisms and biological transformations. Microorganisms responsible for the
microbial degradation of organic micropollutants obtain their energy by degradation and oxidation of
organic carbon. The process of this energy production is primarily based on redox reactions. Energy is
released by electron transfer from an electron donator (organic carbon) to an electron acceptor (O2, NO3
-,
NO2
-, Mn(IV), Fe(III), SO4
2-, CO3
2-) and is then stored within the organism biochemically and utilized for
its cell growth or reproduction. From the pool of potential electron acceptors microbial communities always
prefer those for which the redox reactions provide maximum energy release. The energy gain is highest at
aerobic respiration and lowest at methanogenesis. As a consequence, electron acceptors are successively
consumed, causing a succession of individual redox zones in the underground. This means for the order of
potential electron acceptors, that oxygen is consumed as long as it is available. After its depletion, oxygen is
followed up by NO3
-, Mn(IV), Fe(III), SO4
2- and finally by CO3
2- being transformed to methane. Depending
on the thickness of the surface sediments, the quality of the infiltrating water, and the hydrochemical,
biological and hydraulic boundary conditions of the aquifer, the succession of the redox zones can occur on
a flow distance of a few centimeters up to several 10 m. Certain microorganisms can perform their
microbial activity only at certain redox conditions implying that the microbial community is variable in the
197
different redox zones. Furthermore, the extent of biodegradation of an organic micropollutant is dependent
on sufficient residence times and sometimes proceeds retarded. Thus, the degradation of organic
micropollutants is linked to both the presence of favorable redox conditions and sufficient residence times
in these zones. As a consequence, elimination rates of compounds intrinsically degradable in distinct redox
zones may vary considerably depending on local geological and hydrochemical conditions as well as on
organic loads of surface waters and infiltration zones.
Characteristics of the bank filtrate are often affected by changes of the surface water quality that is
characterized by the number of particles, concentration of dissolved organic matter from natural and
artificial sources, oxygen, ammonia, nutrients, microorganisms, and other pollutants. The Rhine river is an
excellent example how changes in surface water quality influence the characteristics of the corresponding
bank filtrate. Figure 1 summarizes the concentrations of ammonia, manganese, and oxygen in the bank
filtrate of the Rhine over a period of several years. In the early 1970s, Rhine water was highly polluted.
Ammonia was present and nearly no oxygen. Due to the reduction of biodegradable organic material during
infiltration the little oxygen present in the surface water and even nitrate were consumed, the aquifer was
characterized by an anaerobic redox status, in which iron and manganese were reduced and released from
the soil. In the mid 1980s, Rhine water quality improved because of better municipal and industrial
wastewater treatment and its oxygen concentrations increased. As a consequence, conditions in the aquifer
became aerobic, iron and manganese stayed in the insoluble oxidized form and finally disappeared from the
bank filtrate [1].
Fig. 1. Development of bank filtrate quality at the Rhine river.
Since redox conditions of an aquifer may change over time and from one aquifer to another, it is important
to clarify the potential impact of such redox shifts on the removal capacity of riverbank filtration sites with
regard to organic micropollutants commonly present in surface waters. Therefore, the present study was
performed to investigate the fate of selected organic micropollutants during riverbank filtration at sites that
are characterized by different boundary conditions, in particular varying redox conditions (aerobic, aerobic-
denitrifying, denitrifying, strictly anaerobic) and travel times of the infiltrated water in the underground
(5-300 days).
Selection of Target Compounds
The selection of individual organic micropollutants to be investigated was carried out under consideration
of their ubiquitous application and spreading, their detectability in German surface waters and their
physico-chemical properties that should be as diverse as possible. Taking these criteria into account, the
following substance classes and single compounds were selected. All target compounds are solely of
anthropogenic origin.
0.0
0.5
1.0
1.5
1973
1975
1977
1979
1981
1983
1985
1987
1989
1991
1993
1995
1997
ammonia, manganese [mg/L]
0.0
1.0
2.0
3.0
4.0
oxygen [mg/L]
ammonia
oxygen
manganese source: ARW (1998)
198
Aminopolycarboxylates
In several industrial and domestic processes unwanted metal ions cause trouble. In such cases chelating
agents like aminopolycarboxylates can be applied that form stable and water-soluble complexes with
multivalent metal ions and restrict them from playing their normal chemical role. The global consumption
of aminopolycarboxylates in various fields of industry and private households amounts to roughly
200,000 t/a, with an increasing tendency. Important fields of application comprise water softening and
treatment, pulp and paper production, agro chemicals, detergents in industry and private households, food
processing, galvano industry, textiles, cosmetics, pharmaceuticals, and medical detoxification.
Quantitatively prominent representatives of this substance class are ethylenediaminetetraacetic acid
(EDTA), nitrilotriacetic acid (NTA), and at a progressive rate diethylenetriaminepentaacetic acid (DTPA).
Due to their wide fields of industrial and domestic application and their high polarity,
aminopolycarboxylates enter our creeks, rivers, and lakes, mainly via industrial and domestic waste waters
and can be detected in many European rivers at considerable concentrations in the µg/L range [4].
Aromatic Sulfonates
Among other things, aromatic sulfonates have been used for more then 100 years as important building
blocks in the production of dyestuffs. Further important application fields comprise optical brighteners,
synthetic leather tanning agents, textile auxiliaries, paper chemicals, concrete additives, and intermediates
in the manufacture of pharmaceuticals. Due to their complete dissociation in water, aromatic sulfonates are
highly soluble in water and enter the aquatic environment via industrial and domestic waste waters. Besides
this, their release from concrete liquefiers used on construction sites as well as from building rubble
recycling material (naphthalene-formaldehyde condensates) is of certain importance. In running waters in
particular substituted naphthalene sulfonates occur. Important representatives are 2-aminonaphthalene-1,5-
disulfonate, naphthalene-1,3,6-trisulfonate, napthalene-1,5-disulfonate, napthalene-1-sulfonate, and
napthalene-2-sulfonate. Typical surface water concentrations of these single compounds are in the µg/L
range, however, at some sites even peak concentrations in the mg/L-range were found [5].
Pharmaceutical Compounds
Drugs are prescribed and used for treatment and therapy of a multitude of human diseases. Alone in
Germany, the annual consumption volumes of the most frequently used medicaments range often between
10 and 200 tons. Following oral intake, many of the applied pharmaceutically active substances are
absorbed only incompletely from the gastro-intestinal tract. The non-absorbed fraction is normally excreted
directly, whereas the absorbed fraction leaves the body often in a modified form, e.g. as glucuronide
conjugate following metabolic conversion. The excretory products reach the sewage treatment plants, where
they are not eliminated in the course of conventional treatment and finally are discharged into the rivers.
Conjugates formed within the human body are often cleaved during wastewater treatment, releasing the
pharmaceutically active compound originally applied. Effluents of sewage treatment plants often contain
pharmaceuticals at concentrations between 0.01 µg/L and more than 1µg/L. In receiving surface waters,
concentrations between a few ng/L and several µg/L are detected, depending on the respective waste water
portion. Additive levels of all detectable pharmaceutical compounds amount often to more than 1 µg/L.
Typical and often found representatives of this substance class are diclofenac (antiphlogistic),
carbamazepine (antiepileptic), bezafibrate (lipid-regulator), and sulfamethoxazol (antibiotic) [6,7].
Iodinated X-Ray Contrast Media
This substance class comprises various triiodobenzene derivatives. Following application, these compounds
can make organs visible that are in other respects not depictable. The contrast medium in the tissue to be
investigated absorbs X-rays more strongly than vicinal body tissues providing a sufficiently strong contrast
for radiographic analysis. In typical X-ray examinations, these substances are administered in doses of up to
100 g per patient and application. The global application volume of iodinated X-ray contrast media in the
field of medical diagnosis amounts to about 3,500 tons per year. Iodinated X-ray contrast media are highly
water soluble and stable with the purpose of being metabolically resistant and rapidly excreted from the
body (~ 95 % within 24 hours). As a consequence, following administration, they are excreted into waste
waters via urine and reach the sewage treatment plants, where they are typically not eliminated. In several
sewage treatment plant effluents concentrations of 10 µg/L and more were found. The typical pollution of
199
German surface waters ranges from 1 ng/L to 1 µg/L. Typical iodinated X-ray contrast media are iomeprol,
amidotrizoic acid and iopamidol [8].
MTBE
Methyl tertiary-butyl ether (MTBE) that is added to fuel as anti-knocking agent is among the most produced
mass chemicals in the US. The addition of MTBE to fuel (up to 15 % by weight) prevents an uncontrolled
combustion of the fuel/air-mixture by providing the oxygen necessary for a complete combustion. MTBE is
volatile and well water soluble and enters the aquatic environment presumably on diffuse pathways, in
particular via air pollution. Direct emissions by leakages and losses of fuel in the course of production,
storage and application are certainly also of major importance. Typical MTBE concentrations in German
surface waters were found to be in the range of 0.2-0.3 µg/L [9].
Investigations in the Field
So far, data from systematic studies investigating the impact of residence times and redox conditions on the
fate of the mentioned organic micropollutants during riverbank filtration is not available. For investigation
of these boundary conditions extensive measurement campaigns were carried out at four well-characterized
aquifers along the rivers Rhine, Ruhr and Elbe. Figure 2 shows the particular cross section of each
riverbank filtration site studied.
Fig. 2. Cross-section profiles of the investigated aquifers.
Testing Field – Rhine A
Along the Rhine, several water works extract and treat Rhine bank filtrate for drinking water production.
The water works presented here is located in the lower Rhine area. At this site, extraction wells are situated
at a distance of about 170 m from the bank in parallel to the river. The wells extract a mixture of Rhine
bank-filtered water and natural land-sided groundwater from the sandy-gravely, about 15 m thick terrace
deposits of the Rhine. The aquifer is characterized by a pronounced anisotropy of hydraulic conductivities.
Thereby, the lower layer shows an at least six-fold higher permeability than the upper layer. Due to an
almost impermeable layer at the river bank, the water infiltrates in particular at the river bottom in the
P
A
1
X
Y
Z
1
1
1
2
Rhine A Rhine B
Ruhr
Elbe
collection well
Rhine
Rhine
collection well
collection well
reservoir
200
middle of the river. The majority of the water collected by the wells originates from the lower aquifer layer
and the produced raw water contains about 60 % Rhine bank filtrate. Redox conditions are predominantly
aerobic. Between well gallery and bank line several sampling wells have been constructed that are in
contrast to the collection wells not influenced by natural land-sided groundwater. Investigations presented
here were conducted at the sampling well A1 with deep filter position. The infiltrated river water needs
between 7 and 20 days to pass this well (residence time in the aquifer). Fortunately, this site was
investigated intensively already in previous projects, providing an impressive data set covering a period of
several months.
Testing Field – Rhine B
Also this site is located at the right bank of the lower Rhine forming a gallery of several vertical wells
parallel to the bank line situated approximately 60 m from the bank. Again, the infiltration into the clay and
fine-sand layers of the 10-12 m thick aquifer proceeds mainly at the river bottom. Between Rhine and
collection wells several sampling wells are available, that are typically not influenced by natural land-sided
groundwater. The investigated sampling point P provides three independent sampling wells with three
different filter depths. Travel times of the infiltrated water to pass the sampling well are approximately
60 days for well P1 (lower layer), 30 days for well P2 (middle layer) and 12 days for well P3 (upper layer).
The water extracted by the collection wells is a mixture water of various age and contains approximately a
portion of 25-30 % natural land-sided groundwater. Samples were taken in summer 2003 at 14 dates from
the Rhine and at 5 dates from the three sampling wells (P1, P2, P3). During the time of sampling in summer
2003, redox conditions were at the edge of aerobic to denitrifying.
Testing Field – Ruhr
This study site is a pure testing field situated at the middle course of the Ruhr river. In this area the Ruhr
river is dammed in a branch. Due to a retaining weir upstream and the controlled outflow of the reservoir a
very constant reservoir water level can be maintained forming a potential gradient of more than 5 m
between the surface water and the adjacent about 5 m thick, sandy-gravely and well permeable aquifer. The
investigated bank filtrate moves without any natural groundwater influx from the infiltration point at the
reservoir bottom along the hydraulic gradient to the water downstream of the reservoir retaining weir. This
aquifer was shown to be strictly anaerobic since investigations have demonstrated a completed sulfate
reduction already in the infiltration zone. Along the aquifer several sampling wells are available. Travel
times are comparatively short and constitute between 5 and 15 days depending on the location of the
sampling point. Samples from this testing field were taken in the period from autumn 2002 to summer 2003
at six dates. At each date samples from the surface (reservoir) water and from several sampling wells were
investigated.
Testing Field – Elbe
This study site is located at the middle course of the Elbe river at its left bank. The water extracting plant
used by the local water works comprises of well galleries with several vertical wells situated at a distance of
about 300 m from the bank in parallel to the Elbe river. The thickness of the aquifer is about 40-55 m and is
characterized by a certain anisotropy of hydraulic conductivities causing intensive mixture effects between
river and collection wells. The aquifer can be described as a three-layer system. The middle layer is formed
by medium sand with a hydraulic conductivity of about 6 ּ 10-4 m/s. The upper and lower layers are formed
by coarse sand to gravel, with a hydraulic conductivity of 2 ּ 10-3 m/s. Due to silt bands in the lower
section of the aquifer an undercurrent of the river with groundwater from the catchment area right-sided to
the Elbe occurs. The water extracted by the collection wells contains a portion of approximately 60 %
riverbank filtrate. Between well gallery and Elbe bank several sampling wells exist that are not influenced
by any land-sided natural groundwater influx. These bank filtrate sampling points were constructed as
groups of three sampling wells of different depth. Retention times of the infiltrated water in this aquifer are
comparatively long. Within the middle sand layer section, the water needs more than 150 days to reach
sampling well Z2. Travel times within the better permeable upper layer (sampling wells X1, Y1, Z1) are
between 45 and 80 days. Investigations concerning the redox status have demonstrated denitrifying
conditions already in the infiltration zone. Samples from this testing field have been taken so far only once
(in April 2003). Samples were taken from the surface water, from three sampling wells of the upper aquifer
layer, and from one sampling well of the middle layer.
201
Tab. 1. Characteristics of the investigated riverbank filtration sites
Rhine A Rhine B Elbe Ruhr
aquifer thickness 12-15 m 10-12 m 40-55 m 5 m
material gravel/sand gravel/sand gravel/sand gravel/sand
residence time 7-20 days 12-60 days 45-300 days 5-15 days
Kf-value 12 • 10-3 m/s
(deep, A1)
2 • 10-3 m/s
(high)
3,0 • 10-3 m/s
to
6,0 • 10-3 m/s
0,6 • 10-3 m/s
(depth 2)
2,0 • 10-3 m/s
(depth 1)
10-1 m/s
to
10-3 m/s
redox conditions aerobic aerobic to
denitrifying
denitrifying strictly anaerobic
As becomes apparent from table 1, investigations at these well characterized bank filtration sites allow,
based on a comprehensive data evaluation, conclusions to be drawn concerning the capacity of riverbank
filtration for elimination of organic micropollutants at different travel times and redox conditions.
Results of Investigation
Aminopolycarboxylates
Figure 3 shows the results of the studies at the specified riverbank filtration sites with their site-specific
characteristics and boundary conditions for EDTA. As can be seen from the diagrams, EDTA can be
regularly detected in the bank filtrates of the different sites, even though concentrations are in most cases
slightly lower than in the corresponding surface water. The impact of travel times and pathway lengths on
EDTA elimination seems to be rather low. With regard to the impact of redox conditions it can be
postulated, that the reduction in EDTA concentrations under aerobic conditions is slightly more pronounced
than under denitrifying and anaerobic redox conditions.
Fig. 3. EDTA concentrations in surface waters and bank filtrates.
In the bank filtrate samples from the testing field at the Ruhr surprisingly even higher EDTA concentrations
were detected than in the corresponding reservoir surface water. This finding can be ascribed most probably
Rh
ein
P
1
18. Jun
20. Jun
25. Jun
30. Jun
9. Jul
14. Jul
18. Jul
23. Jul
28. Jul
1. Aug
6. Aug
11. Aug
15. Aug
20
. Aug
0
2
4
6
8
10
c in µg/L
Elbe
X1
Y1
Z1
Z3
Apr 2003
0,0
2,0
4,0
6,0
c in µg/L
Ruhr
I/A - Neu 1
I/A - B 2
I/B 3
I/B - C 2
I/C 2
I/C - R 2
Okt 2
De
z
M
0,0
2,0
4,0
6,0
8,0
10,0
c in µg/L
Jan 94
Mrz 94
Mai 94
Jul 94
Sep 94
Nov 94
Jul 96
Sep 96
Nov 96
Jan 97
Mrz 97
Mai 97
Jul 97
Sep 97
Nov 97
Jan 98
Mrz 98
Mai 98
Jul 98
Rhein
0
4
8
12
16
c in µg/L
spot samples 1994-1998
spot samples Jun-Aug 2003
spot samples
Okt 2002 - Jun 2003
Rhine
A1
Rhine
P3
P2
P1
Rhine A
Rhine B
Elbe
Ruhr
bank filtrate
river
202
to the complexing agent DTPA that occurs in the Ruhr river at relatively high concentrations. According to
Stumpf et al. [10], DTPA can be partially degraded by microorganisms generating the aminopoly-
carboxylates NTA and EDTA as metabolites.
In order to investigate this aspect more precisely, figure 4 shows the progression of the molar
aminopolycarboxylate concentrations (NTA, EDTA and DTPA) along the aquifer at the Ruhr. Thereby,
mean values were calculated for each sampling point from all measurements performed. The coincidence of
DTPA degradation and increase of EDTA concentrations becomes quite obvious. Overall, DTPA is reduced
at an intermediate degradation rate while EDTA is only slowly degraded along the aquifer. The reason for
observing elevated EDTA concentrations only in the bank filtrate at the Ruhr is most probably due to the
fact that DTPA concentrations in Rhine and Elbe are comparatively lower than in the Ruhr, so that such an
effect proportionally cannot become noticeable in these rivers. In riverbank filtrates of Rhine and Elbe
DTPA could only be sporadically detected. Even in surface waters, DTPA concentrations are often close to
the analytical quantitation limit making an evaluation of the results generally difficult. However, DTPA
seems to be slightly better degradable than EDTA. The complexing agent NTA can be easily degraded by
microorganisms and its elimination turned out to be independent from redox conditions and travel times.
According to the results of the present study, the entire elimination of NTA proceeds already in the first few
centimeters to meters of the particular underground passage.
Fig. 4. Molar aminopolycarboxylate concentrations along the aquifer at the Ruhr and postulated
degradation of DTPA [10].
Aromatic Sulfonates
According to the findings of the studies presented here, individual naphthalene sulfonates show major
differences with regard to their elimination during riverbank filtration, depending on their molecular
structure and substitution rate. However, for the individual naphthalene sulfonates discussed here, the
particular extent of elimination turned out to be rather independent from the prevailing redox potential.
While the two singly substituted compounds naphthalene-1-sulfonate and naphthalene-2-sulfonate are well
eliminated during riverbank filtration, representatives with a higher substitution rate like
2-aminonaphthalene-1,5-disulfonate and naphthalene-1,3,6-trisulfonate show a rather persistent behavior.
Independently from the travel time in the underground and from the underlying redox conditions, no
decrease in concentration was observable for 2-aminonaphthalene-1,5-disulfonate. This finding emphasizes
the fact that the analyzed riverbank filtrate samples were in the present studies indeed free of any major
dilution by natural land-sided groundwater. In contrast to 2-aminonaphthalene-1,5-disulfonate, the
elimination extent of naphthalene-1,3,6-trisulfonate was significantly dependent on the residence time of
the water in the underground. This was observable not only within the testing fields Rhine B and Ruhr, but
also by comparison of the aquifers among one another (figure 5). In particular the aquifers Elbe and
Rhine B with their comparatively longer retention times showed a more pronounced decrease in
naphthalene-1,3,6-trisulfonate concentrations.
N
N
HOOC
HOOC
N
COOH
COOH
COOH
N
COOH
HOOC
HOOC
N
N
COOH
COOH
COOHHOOC
NTA
EDTA
DTPA
Ruhr
I/A - B 2
I/B 3
I/B - C 2
I/C 2
I/C - R 2
c in [nmol/L]
0
5
10
15
20
25
30
35
40
DTPA
EDTA
NTA
203
Fig. 5. Concentrations of naphthalene-1,3,6-trisulfonate in surface waters and bank filtrates.
Pharmaceutical Compounds
The fate of the investigated pharmaceutical compounds during bank filtration has to be evaluated
differentially. Even though the lipid-regulator bezafibrate was permanently present in the rivers Rhine,
Ruhr and Elbe, it was not detected in any sample of the corresponding underground passage. Thus,
bezafibrate is independently from the travel time in the underground and the underlying redox conditions
generally well removable during riverbank filtration. The extent of elimination of the antiphlogistic
diclofenac, the antiepileptic carbamazepine, and the antibiotic sulfamethoxazol was, however, clearly
dependent on the underlying redox milieu. While carbamazepine and sulfamethoxazole were only slightly
removable under aerobic and denitrifying conditions, both substances could be only sporadically detected
(and then at very low concentrations) in the strictly anaerobic bank filtrate at the Ruhr implying their
preferential removal at strictly anaerobic conditions.
Fig. 6. Concentrations of diclofenac in surface waters and bank filtrates.
Apr 96
Mai 96
Jun 96
Jul 96
Aug 96
Sep 96
Okt 96
Nov 96
Dez 96
Jan 97
Feb 97
Mrz 97
Apr 97
Mai 97
Jun 97
Jul 97
Rhein
Uferfiltrat
0,0
0,1
0,2
0,3
c in µg/L
Ruhr
I/A - Neu 1
I/A - B 2
I/B 3
I/B - C 2
I/C 2
I/C - R 2
Okt 02
Nov 02
Dez 02
Apr 03
Mai 0
3
Jun
0
0,0
0,1
0,2
0,3
0,4
c in µg/L
Elbe
X1
Y1
Z1
Z2
Apr 2003
0,0
0,1
0,2
0,3
0,4
c in µg/L
spot samples
Okt 2002 - Jun 2003
spot samples 1996-1997
Rhine
A1
Rhine A Elbe
Ruhr
Rhein
PFB3
PFB2
PFB1
18. Jun
20. Jun
25. Jun
30. Jun
9. Jul
14. Jul
18. Jul
23. Jul
28. Jul
1. Aug
6. Aug
11. Aug
15. Aug
20. Aug
0
0,2
0,4
0,6
0,8
c in µg/L
spot samples Jun-Aug 2003
Rhine
P3
P2
P1
bank filtrate
river
Rhine B
Rhei
n
PF
B3
P
FB
PF
18. Jun
20. Jun
25. Jun
30. Jun
9. Jul
14. Jul
18. Jul
23. Jul
28. Jul
1. Aug
6. Aug
1
1. Aug
5A
ug
Aug
0
25
50
75
100
125
150
c in ng/L
Elbe
X1
Y1
Z1
Z2
Apr 2003
0
20
40
60
80
100
120
c in ng/L
Ruhr
I/A - Neu 1
I/A - B 2
I/B 3
I/B - C 2
I/C 2
I/C - R 2
Okt 02
Dez 02
Mai 03
0
50
100
150
200
250
c in ng/L
Jun
97
Jul
97
Aug
97
Sep
97
Okt
97
Nov
97
Dez
97
Jan
98
Feb
98
Mrz
98
Apr
98
Mai
98
Jun
98
Jul
98
Aug
98
Sep
98
Okt
98
Nov
98
Dez
98
Jan
99
Feb
99
Mrz
99
Apr Mai
Jun
Rhein
Uferfiltrat
0
100
200
300
400
500
600
c in ng/L
Ruhr
Elbe
Rhine A
Rhine B
spot samples 1997-1999
Rhine
A1
spot samples
Oct 2002 - Jun 2003
spot samples Jun-Aug 2003
Rhine
P3
P2
P1
bank filtrate
river
204
On the other hand, diclofenac turned out to be well removable at aerobic and denitrifying redox conditions,
while the results from the Ruhr aquifer imply a significantly lower degradation rate during strictly
anaerobic riverbank filtration (figure 6).
Iodinated X-ray Contrast Media
Also the fate of iodinated X-ray contrast media is clearly dependent on the respective boundary conditions
of an aquifer. Iomeprol turned out to be well eliminated during bank filtration, independently from
residence times and redox conditions. Amidotrizoic acid, however, is not removable at aerobic conditions,
but showed a significant concentration decrease at aquifers providing denitrifying or strictly anaerobic
conditions. Similar results were found for iopamidol. However, in comparison to amidotrizoic acid, the
removal of iopamidol responds much more sensitively to redox changes in the aquifer. While no removal
could be observed for iopamidol at the aerobic aquifer at the Rhine (testing field Rhine A), its concentration
level was already significantly lower in the bank filtrate from the aerobic-denitrifying aquifer (Rhine B). At
this aquifer, the lowest iopamidol concentrations were detected in particular at those dates when the redox
potential was drifted more clearly into the denitrifying region. A good elimination of iopamidol was also
observed at the denitrifying Elbe aquifer and at the strictly anaerobic bank filtration site at the Ruhr. Thus,
iopamidol might represent a quite sensitive redox marker for this kind of studies.
Fig. 7. Concentrations of iopamidol in surface waters and bank filtrates.
MTBE
During aerobic bank filtration at the Rhine (historical data from testing field Rhine B) MTBE was found to
be only insufficiently removed and could be always detected in the raw water of the local water works.
Concerning the fate of MTBE during bank filtration at denitrifying and strictly anaerobic milieu conditions
no statement can be given so far, since MTBE concentrations in surface waters of the relevant testing fields
were often too low and existing data sets are too fragmentary at present.
Conclusions
The systematic investigations at aquifers with geological and hydrochemical boundary conditions presented
here revealed that the extent of elimination of several organic micropollutants is decisively dependent on
the underlying redox conditions. Furthermore, the fate of organic micropollutants during riverbank filtration
and underground passage can hardly be predicted, since chemically relatively similar compounds
Jan 02 Apr 02 Jul 02 Okt 02 Mai 03
Rhein
A1
0
100
200
300
400
500
600
c in ng/L
Rhine A
Elbe
X1
Y1
Z1
Z2
Apr 2003
0
5
10
15
20
25
30
35
c in ng/L
Elbe
Ruhr
I/A - Neu 1
I/A - B 2
I/B 3
I/B - C 2
I/C 2
I/C - R 2
Okt 2002
Nov 20
0
Dez 2
0
A
pr
0
Mai
Ju
0
100
200
300
400
500
600
700
c in ng/L
Ruhr
spot samples
Oct 2002 - Jun 2003
Rhein
PFB2
18. Jun
20. Jun
25. Jun
30. Jun
9. Jul
14. Jul
18. Jul
23. Jul
28. Jul
1. Aug
6. Aug
11. Aug
15. Aug
20. Aug
0
100
200
300
400
500
600
c in ng/L
spot samples Jun-Aug 2003
Rhine
P3
P2
P1
bank filtrate
river
Rhine B
Rhine
205
sometimes show quite different elimination rates. The major part of the elimination proceeds always in the
first few meters of the infiltration pathway. Nonetheless, longer residence times can significantly improve
removal rates of some individual compounds (e.g. naphthalene-1,3,6-trisulphonate). However, it became
also apparent that the underlying redox conditions have a more pronounced effect on the removal capacity
than residence times of the water in the underground.
References
[1] Schmidt CK, Lange FT, Brauch H-J, Kühn W, 2003, Experiences with riverbank filtration and infiltration in Germany:
Proceedings International Symposium on Artificial Recharge of Groundwater, 14.11.2003, Daejon, Korea, 115-141; [2] Ray C, Melin
G, Linsky RB, 2002, Riverbank Filtration: Kluwer Academic Publishers, Dordrecht, The Netherlands; [3] Grischek T, 2003, Zur
Bewirtschaftung von Uferfiltratfassungen an der Elbe: Thesis, Technical University Dresden; [4] Schmidt CK, Brauch H-J, 2003,
Aminopolycarbonsäuren in der aquatischen Umwelt. Quellen, Vorkommen, Umweltverhalten, Toxizitäten und Beseitigung:
Schriftenreihe des DVGW-Technologiezentrums Wasser (TZW) Band 20, Karlsruhe; [5] Lange FT, Redin C, Brauch H-J, Eberle SH,
1998, Auftreten aromatischer Sulfonate in Industrieabwasser, Flusswasser, Uferfiltrat und in der Trinkwasseraufbereitung: Vom
Wasser, 90, 121-134; [6] Heberer T, 2002, Occurrence, fate, and removal of pharmaceutical residues in the aquatic environment: A
review of recent research data: Toxicology Letters, 131, 5-17; [7] Sacher F, Brauch H-J, 2002, Rückstände von Arzneimitteln und
endokrin wirksamen Stoffen in Gewässern: Veröffentlichungen aus dem Technologiezentrum Wasser, 18, 1-12; [8] Ternes TA, Hirsch
R, 2000, Occurrence and behavior of X-ray contrast media in sewag facilities and the aquatic environment, Environmental Science &
Technology, 34, 2741-2748; [9] Achten C, Kolb A, Püttmann W, 2002, Methyl tert-butyl ether (MTBE) in river and wastewater in
Germany: Environmental Science & Technology, 36, 3652-3661; [10] Stumpf M, Ternes TA, Schuppert B, Haberer K, Hoffmann P,
Ortner HM, 1996, Sorption und Abbau von NTA, EDTA und DTPA während der Bodenpassage: Vom Wasser, 86, 157-171.
________________________
Carsten K. Schmidt, Ph.D. (* 25.03.1971). Carsten K. Schmidt started his work at the DVGW-Water Technology Center in
Karlsruhe (TZW), Germany, in January 2002. During the last years, he has been responsible for a bank-filtration project (financially
supported by the German Federal Ministry of Education and Research) concerning the fate of organic micropollutants and
corresponding optimization strategies in dependency on site-specific boundary conditions. His background includes about 10-years of
experience in development and validation of trace analysis methods for the determination of organic micropollutants. Before joining
the Chemical Analysis Department of the TZW, he studied chemistry in Clausthal-Zellerfeld, Hannover (Germany), and Cork (Ireland)
and received a Ph.D. in Analytical Chemistry and Toxicology from the University of Hannover.
